Cadmium exposure via diet and its implication on the ... - Nature

Journal of Exposure Science and Environmental Epidemiology (2015) 25, 433–442 © 2015 Nature America, Inc. All rights reserved 1559-0631/15

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ORIGINAL ARTICLE

Cadmium exposure via diet and its implication on the derivation of health-based soil screening values in China Mao-Sheng Zhong1,2,3, Lin Jiang1,2,3, Dan Han1,2,3, Tian-Xiang Xia1,2,3, Jue-Jun Yao1,2,3, Xiao-Yang Jia1,2,3 and Chao Peng1,2,3 The cadmium (Cd) intake rates via diet of adults from different regions in China were between 0.160 and 0.557 μg/(kg BW·day), which were less than the provisional tolerable monthly intake (0.833 μg/(kg BW·day)) issued by Food and Agriculture Organization/ World Health Organization in 2010, but higher than the one (0.365 μg/(kg BW·day)) issued by the European Food Safety Authority in 2011, to protect children, vegetarians and people living in heavily contaminated regions, and the intake rate of children (1.007 μg/(kg BW·day)) at the national scale was higher than the values recommended by the above institutes and those of adults. Vegetables were the critical contributors, followed by rice, flour, meats and aquatic products. Cd concentration in vegetable was the most sensitive factor in calculating the intake rate, followed by its contents in rice and aquatic products, and the intake rate of flour, indicating that more attention should be given to these parameters in future total diet surveys. When dietary exposure was incorporated, the derived national screening value of Cd under commercial scenario was reduced from 825 to 458 mg/kg, while the values of the north, south, Beijing and Shanghai were reduced to 627, 365, 693 and 489 mg/kg, respectively, indicating that the hazard would be underestimated if dietary exposure was not taken into account, especially for the south. The great variance between the screening values was due to the varied Cd intake rates, which indicated that deriving a screening value for each specific area based on its corresponding exposure characteristics was more appropriate. The national screening level for the residential scenario derived theoretically based on the dietary exposure characteristics of children was a negative value, meaning that the dietary intake rate was above the tolerable value. The method used in the United Kingdom to derive soil guideline values when non-soil exposure accounted for more than half of the maximum tolerable daily intake dose may be an appropriate estimate, but the exact ratio assigned to soil exposure should be assessed comprehensively based on a more sophisticated dietary exposure survey and the corresponding economic implications. Journal of Exposure Science and Environmental Epidemiology (2015) 25, 433–442; doi:10.1038/jes.2015.5; published online 4 March 2015 Keywords: Cd; contaminated sites; dietary exposure; soil screening value

INTRODUCTION Land contamination is a serious problem in China. In recent years, many old and polluting industries are being shut down or relocated away from urban centers due to rapid urban growth in China. As a result, a large number of contaminated sites (often referred to as “brownfields”) are emerging. These brownfields pose a dual problem: an environmental and health hazard in China’s most densely populated areas, and as brownfields can not be developed safely, they pose an obstacle to urban and economic development. Therefore, a health-risk assessment must be carried out and remediation must be implemented before the redevelopment of brownfield sites, in accordance with the provisional rules for the Environmental Management of Contaminated Sites and notes on strengthening pollution control throughout the decommissioning, relocation and redevelopment of industrial enterprises issued by Ministry of Environmental Protection of China and the related administrative rules issued by local environmental protection bureaus.1–6 According to the technical guidelines of China, a tiered assessment approach is recommended.7–9 This approach is also used worldwide and is advantageous in saving investigation costs,

while maintaining the accuracy of results.10,11 The assessment begins with the simple process of taking soil samples, analyzing them for relevant contaminants and comparing the results with their corresponding soil screening values (SSVs), which are defined as the allowable chemical concentrations in environmental media below which no additional regulatory attention is warranted. If the results are below the SSVs, the risk to human health is deemed acceptable and no further investigation is needed. If the results are high, a potential risk to human health is indicated and a more detailed site-specific assessment (often called tier two assessment) is needed, and some form of action or management may be necessary. Therefore, SSVs are very important tools in identifying contaminants of concern to which more resources must be allocated to investigate, and in some countries it is used as a tool to screen out sites with high health risks for which more active management measures should be taken. Although they may be modified according to the analysis detection limits of contaminants or background concentrations found naturally, most SSVs are health-based and are derived using a risk-assessment approach, combining toxic potency estimates, acceptable target risks and hazards, and default conservative

1 Beijing Municipal Research Institute of Environmental Protection, Beijing, China; 2National Engineering Research Centre of Urban Environmental Pollution Control, Beijing, China and 3Beijing Key Laboratory for Risk Modeling and Remediation of Contaminated Sites, Beijing, China. Correspondence: Professor Lin Jiang, Beijing Municipal Research Institute of Environmental Protection, No.59 Beiyingfangzhongjie, Xicheng District, Beijing, China. Tel.: +86 010 88362281. Fax: +86 010 88362281. E-mail: [email protected] Received 16 September 2014; revised 19 November 2014; accepted 23 November 2014; published online 4 March 2015

Cadmium exposure via diet Zhong et al

434 exposure values. For carcinogenic contaminants screening values are derived assuming the acceptable individual excess cancer risks from exposure to contaminated soils to be 1 × 10 − 6, whereas for non-carcinogenic pollutants the acceptable hazard indexes are always set to 1, meaning that the exposure dose of the individual contaminant from soil should not exceed its daily tolerable intake. However, the daily tolerable intakes are the maximal allowable doses that receptors can take from all potential exposure pathways. Therefore, hazards from non-carcinogens may be underestimated by applying SSVs derived from the assumption that the acceptable hazard index is 1 from soil exposure if there are some other important exposure pathways. Cadmium (Cd) is a toxic heavy metal that is ubiquitous in environmental media, especially in soils. Concentrations of Cd in soils from smelting plants have been reported to be between 6.35 and 117.9 mg/kg,12–19 which are much higher than the national background values of 0.02–0.33 mg/kg.20 More than 1.3 × 105 km2 of soils are contaminated by Cd in China and 1.46 × 108 kg of agricultural products are polluted by Cd every year.21 Currently, the SSV of Cd is determined by its non-carcinogenic toxicity and derived by assuming the acceptable hazard index to be 1. However, apart from soil exposure, dietary intake is also a critical pathway.22 It was reported that Cd intake from diet was 2.5 μg/kg body weight per week in Shanghai,23 which contributed to 35.7% of the provisional tolerable monthly intake (PTMI) of Cd (7 μg/kg body weight per week) issued by the World Health Organization (WHO) in 2010 (ref. 24). According to Huang et al.,25 Cd exposures through diet in three cities of the Zhejiang Province during 2009 and 2010 were between 6.4 μg/day and 11.4 μg/day, contributing 15.6–42.6% to the PTMI. The result from a dietary survey carried out in 2000 revealed that the national average dietary Cd exposure was 22.2 μg/day,26 which was equivalent to 44.4% of the PTMI. Therefore, if the screening values derived under the current assumptions are applied in a tiered one risk assessment, many Cd-contaminated sites may be screened out as “clean sites” with negligible hazards and the health of future users will be threatened. At present, there are still no national SSVs in China, although the national risk-assessment guidelines have come into effect. The screening values for non-carcinogens issued by the Beijing Environmental Protection Bureau (BEPB) were derived assuming an acceptable hazard index of 0.2 (ref. 27), which was in accordance with some states in the United States. The reason for the conservative assumptions is that receptors may be exposed to five contaminants doing harm to the same organ simultaneously on a contaminated site instead of dietary exposure. Therefore, the aim of this study was to quantify, at the national and regional levels, the average dietary Cd intake by compiling a database of Cd in Chinese foods and the dietary habits of people in different regions. Furthermore, the impact of dietary Cd exposure on SSVs under residential and commercial scenarios was quantified. MATERIALS AND METHODS Food Consumption Data Trends in regional and national dietary habits were obtained from the 2002 China National Nutrition and Health Survey (CNNHS).28 This database is China’s principal nutrition reference, containing the dietary patterns of 68,962 individuals from 31 provinces (excluding Hong Kong, Taiwan and Macau). In the present study, the food types are categorized into one of nine groups: rice and its products (hereinafter referred as rice), wheat flour and other grain products (hereinafter referred as flour), pulses, vegetables, fruits, meats, milk, eggs and aquatic products. Given the diversified dietary habits of different regions, the food consumption patterns were analyzed at national, northern and southern levels. Classification of “north” and “south” was achieved using a geographical divide running from the Huai River to the Qinling Mountains.29 Furthermore, the dietary patterns in Beijing and Shanghai, the most developed and largest cities in China, were

also analyzed. The food consumption rate of the whole nation was directly obtained from the CNNHS survey report,28 while those of the sub-regions (north, south, Beijing and Shanghai) were calculated by weighting the population proportions of the involved provinces (Supplementary Tables S1–S3). For example, the food consumption of the north region was obtained using the following formula: food consumption rate = ΣR × P, where R is the food consumption rate of the provinces in the north region and P is the population proportion of the provinces in the north region.30 Owing to the lack of original diet consumption data for children aged 2–7 years at the provincial level, only the dietary pattern at the national level was compiled based on the age-weighted average values of boys and girls (Supplementary Table S4).

Cd in Different Foods The database of Cd concentrations in different food types was compiled from the published literature since 2000 (Supplementary Tables S5). A total of 22,421 data points for Cd concentrations were collected, but a portion of the data (3,726), which exceeded the standards for contaminants in food,31 was removed with the aim of preventing the negative effect of substandard foods in quantifying the impact of Cd dietary exposure on SSVs. Owing to the lack of original data, concentrations reported in the literatures for different food types were averaged and then used for Cd exposure. As a portion of the data featuring Cd concentrations in rice was obtained from the literature and based on dry weight, and because rice consumption rate is based on fresh weight, the concentrations were converted from dry weight into equivalent fresh-weight concentrations, with the assumption that the water content of rice grain is 10%.32

Calculation of Cd Daily Intake Rate The estimated daily intake of Cd was calculated by multiplying daily food consumption rates with corresponding Cd concentrations according to equation (1) (ref. 30): X EDIDCd ¼ ci ´ IRDi =BW ð1Þ EDIDCd is the estimated daily intake of Cd via diet, μg/(kg BW day); ci is the average Cd concentration in food i, mg/kg; IRDi is the daily intake rate of food i, g/day; and BW is body weight, kg. The reported date for each food type in all literatures were averaged and used to calculate the corresponding EDIDCd at the national level, while for EDIDCd of the south, north, Beijing and Shanghai the statistics reported for the corresponding sub-region were used. Only exposure to the food studied in the present assessment was considered, which account for ~ 85% of the total daily diet intake,28 while the contribution of some food types such as oil, salt, sugar and pastry was excluded, as their contribution to dietary Cd exposure was considered negligible.22,30 Owing to limited data on intake rate and concentration of vegetable in each sub-category (e.g. root and tuber vegetables, leafy vegetable, bulb vegetable etc), Cd exposure from vegetable consumption was calculated based on the averaged concentration of all reported data in literatures multiplied by the total vegetable intake rate recorded in CNNHS, and the variance in concentration and intake rate of the individual sub-category was neglected. For the meat category, the intake rate includes pork, beef, mutton, chicken and duck, and the corresponding concentration was the average value of all reported data on each sub-category. Edible offal (e.g., kidney, liver, etc), always considered to be of high Cd concentration, was not considered in the present assessment, in order to avoid the uncertainty due to limited available concentration data and less intake frequency (4.7 g/day for adults while 3.2 g/day for children according to Jin28). Khan et al.33 and Gemma et al.34 reported that cooking practices did not reduce the concentration of Cd in food. Therefore, the reported concentrations on raw food in the literatures were assumed to be equivalent to the ingested concentrations in the present study.

Derivation of SSV SSVs for residential and commercial exposure scenarios were derived in accordance with the national risk-assessment guidelines of China. Children younger than 7 years of age were the critical receptor for the residential scenario, while adults were the critical receptor for the commercial scenario. Owing to the low volatility of Cd and the negligible hazard from the particulate inhalation pathway, only incidental oral ingestion and dermal adsorption were used as predominant exposure pathways in

Journal of Exposure Science and Environmental Epidemiology (2015), 433 – 442

© 2015 Nature America, Inc.


435 Table 1.

Definition and values of parameters

Parameters

Definition

Values Residential

Reference dose, μg/(kg BW·day) Soil ingestion rate, mg/day Exposure frequency, day/a Skin exposure area, cm2 Soil adherence factor, mg/cm2 Dermal adsorption factor, unitless Body weight, kg

RfD IRs EF AR AH AF BW

1

ðRfD - EDIDCd Þ ´ BW ´ 365 IRs ´ EF

1 þ ðRf D - EDI

1

DCd Þ ´ BW ´ 365 AR ´ AH ´ AF ´ EF

´ 103

Commercial 1

200 350 2448 0.2

100 250 2856 0.2 0.001

15.9

deriving the SSV.11,35 Based on the above assumptions, equation (2) (ref. 36) was used to derive the SSV of Cd. SSV ¼

Sources

ð2Þ

56.8

37 2 2 2 2 2 2

concentration, and the differences between pulses and meats were not obvious. Cd concentrations in each food group in Beijing and Shanghai were lower than the corresponding average values of the whole nation, the south and the north.

RESULTS Chinese Dietary Patterns As shown in Figure 1, the diet intake rates and patterns of the whole population and each sub-region are different. For adults, the total diet intake rates of Beijing and Shanghai are 1066.1 and 1150.2 g/day, respectively, which are higher than the average values of the whole population (874.3 g/day), the north (779 g/ day) and the south (956.2 g/day). Vegetables, rice and flour are the three major food groups, contributing 74.9% and 75.3% of the total diet for the whole population and people of the north, respectively. However, in the south, the dominant components are vegetables, rice and meat, contributing 79.2%. The diet patterns in Beijing and Shanghai, the most developed cities in China located in the north and the south, respectively, are slightly different from those of the whole population, the north and the south. In Beijing, in addition to vegetables, rice and flour, which are 60.5% of the total diet, milk and meat also account for a large percentage, up to 21%. In Shanghai, 23.5% of the total diet is composed of meat and aquatic products, and half is composed of vegetables and rice. The diet intake rate of children at the national level is 589.2 g/day and the pattern is similar to that of adults.

Estimated Daily Cd Intake Based on the national nutrition and health survey in China, the average daily intake for Cd was calculated and the data are listed in Table 2. There was a large variation in Cd intake between regions. Adults in the south had the highest Cd intake rate, up to 0.557 μg/(kg BW·day), whereas adults from the north had 0.240 μg/(kg BW·day). For the adults of the whole nation, the Cd intake rate was 0.444 μg/(kg BW·day), which was higher than that in the north but lower than that in the south. Furthermore, the Cd intake rate of people in Beijing was much lower than that of people in Shanghai, but they were both lower than the national average value and the values of people in the north and south. Compared with adults, the intake rate of children of the whole nation was more than two times higher due to lower body weights. The breakdown of Cd contributions from different food groups also varied between regions. In general, vegetables were the main contributor of Cd intake, while the contributions from pulses, eggs, milk and fruits (except Beijing) were negligible. For the total population, vegetables contributed 27.8% to dietary Cd exposure, followed by rice and flour (Figure 2). For the population in the north, 34.6% was attributed to vegetables, which was comparable to that of the whole population. Flour and aquatic products were also of great importance, accounting for 27.4% and 13.4%, respectively. In contrast, vegetables contributed 34.4% of the total Cd for the south, followed by rice and aquatic products. The main contributors were different between Beijing and Shanghai. For people in Beijing, the most important contributor was vegetables, followed by flour, rice and fruit, but vegetables, rice, meat and aquatic products were the most important contributors for people in Shanghai. The main contributors to the Cd intake rate of children were similar to those of adults due to similar dietary patterns.

Cd in Different Food Groups A total of 18,695 data points of Cd in Chinese foods that were below the standards of contaminants in food were compiled based on the published literature, and the average concentrations are listed in Table 2. The highest mean concentration was found in aquatic products throughout the country, even up to 0.093 mg/kg in the south, followed by the concentrations in rice, flour, pulse, meat and vegetables, and the lowest mean concentration was found in fruits. Furthermore, the concentrations in rice, flour, milk and vegetables from the south were higher than those in the north and the whole country, while Cd in eggs and fruits were lower. Aquatic products from the south had a slightly lower

Screening Value Screening values for commercial and residential scenarios incorporating dietary exposure of Cd were derived and the results are listed in Table 3. When dietary Cd exposure was considered, its corresponding screening values would be reduced. The impact under the commercial scenario is greatest to the south, while it is negligible in the north and Beijing. For the residential scenario, the screening level derived theoretically with dietary exposure incorporated was a negative value, meaning that the dietary Cd intake rate of children, the sensitive receptors under the scenario, exceeded the maximum tolerable dose.

The definition and values of the parameters in equation (2) are summarized in Table 1.

Sensitivity Analysis The sensitivity of the variables (ci and IRDi) used to calculate dietary Cd intake was assessed by calculating the parameters between each input and output during simulations and then evaluating each input’s contribution to the output variance by normalizing to 100%. The Monte Carlo simulation and sensitivity analysis were conducted using @Risk5.5 (Palisade, Palisade Corporation, New York, USA. 2010) software.




436 Sensitivity Analysis As shown in Figure 3, Cd concentration in vegetables was the most sensitive variable in estimating dietary Cd intake. For the

Figure 1.

whole population, the Cd concentration in vegetables contributed ~ 20% to the variance of Cd intake rate, followed by the flour intake rate and Cd concentration in rice and flour. For the people

Dietary patterns of people in different regions.




437 Table 2.

Average concentrations of Cd and intake rates

Food type

Rice Flour Pulses Meat Egg Aquatic products Milk Vegetables Fruit Total

Cd concentration (mg/kg)

Cd intake rate (μg/(kg BW·day))

NL

NH

SO

BJ

SH

NL

NH

SH

BJ

SH

NL-C

0.025 0.043 0.024 0.026 0.008 0.084 0.006 0.028 0.006

0.011 0.015 0.022 0.023 0.011 0.093 0.002 0.021 0.007

0.031 0.053 0.025 0.027 0.006 0.080 0.009 0.035 0.005 /

0.010 0.009 0.002 0.003 0.003 0.019 0.001 0.011 0.013

0.021 0.014 0.013 0.018 0.005 0.043 0.001 0.025 0.001

0.105 0.106 0.007 0.036 0.003 0.044 0.003 0.135 0.005 0.444 (25.2)

0.022 0.066 0.005 0.019 0.006 0.032 0.001 0.083 0.006 0.240 (13.6)

0.180 0.051 0.009 0.051 0.002 0.064 0.004 0.192 0.004 0.557 (31.6)

0.023 0.031 0.001 0.006 0.002 0.009 0.002 0.062 0.024 0.160 (9.1)

0.099 0.014 0.005 0.044 0.004 0.101 0.002 0.138 0.001 0.408 (23.2)

0.246 0.226 0.015 0.089 0.010 0.095 0.012 0.296 0.018 1.007 (16.0)

Abbreviations: BJ, Beijing; NH, north; NL, national level; NL-C, national level of children; SO, south; SH, Shanghai. Values in parentheses are total Cd in take rates in μg/day.

in the north, except for the Cd concentration in vegetables, flour intake rate, Cd concentration in aquatic products, vegetable intake rate and Cd concentration in rice were also very sensitive variables. In contrast, in addition to the Cd concentration in vegetables, the sensitivity of the Cd concentration in rice and aquatic products, and the vegetable intake rate were sensitive variables for the people in the south. For the people in Beijing and Shanghai, in addition to the Cd concentration in vegetables, the Cd concentrations in aquatic products and meat were important. Given the sensitivity of the above variables to the estimated Cd intake via diet, the more accurately their statistic distributions are characterized in future total dietary surveys, the less prone to errors the derived screening values will be. DISCUSSION Through the collection of literature data and based on the national nutrition survey, the present study provides a framework for assessing dietary Cd intake and its impacts on the SSVs in China for the first time. The published data revealed that the Cd concentration of rice from the south, Shanghai and at the national level are higher than the values reported by Canada (0.002 ~ 0.009 mg/kg), Norway (0.013 mg/kg), Portugal (0.015 mg/ kg) and Switzerland (0.011 mg/kg),38–40 lower than the value reported in Japan (0.05 mg/kg) and Iran (0.4 mg/kg),41,42 and similar in comparison with those values reported in Korea (0.02 mg/kg), France (0.024 mg/kg), Finland (0.022 mg/kg), Germany (0.029 mg/kg), the Netherlands (0.029 mg/kg) and the United Kingdom.38–40,43 The differences in Cd concentrations in rice between the north and Beijing are negligible, and they are both similar to the values reported in the previously mentioned foreign countries. Cd concentration in flour from the south and its concentration at the national level are higher than the values reported in Japan, while for the north and Beijing, the concentration in flour is similar to that of Japan.41 The Cd concentrations in meat from the north, south and Shanghai, and its concentration at the national level are similar to the values reported in Japan (0.02 mg/kg) and Belgium (0.024 mg/kg); higher than values reported in Canada (0.00022 ~ 0.004 mg/kg), Denmark (0.0022 mg/kg), Finland (0.001 ~ 0.004 mg/kg), France (0.004 mg/ kg), Germany (0.007 ~ 0.016 mg/kg), Greece (0.0074 mg/kg), the United Kingdom (0.0046 mg/kg and Korea (0.012 mg/kg); and lower than values reported in the Netherlands (0.05 mg/kg) and Norway (0.046 mg/kg).38–40,43 The Cd concentration in meat in Beijing is relatively low and is similar to the values reported in Canada, Denmark, Finland, France, Germany, Greece and the United Kingdom. Except for aquatic products in Beijing, the Cd © 2015 Nature America, Inc.

concentrations in the other regions and the concentration at the national level are higher than the values reported in Finland (0.006 mg/kg), France (0.007 mg/kg), Germany (0.011 ~ 0.116 mg/ kg), Ireland (0.03 mg/kg), Italy (0.0033 mg/kg), the Netherlands (0.01 mg/kg), Norway (0.001 ~ 0.05 mg/kg), Portugal (0.025 mg/kg), Switzerland (0.004 ~ 0.034 mg/kg), the United Kingdom (0.013 mg/ kg), Canada (0.0005 ~ 0.008 mg/kg), Japan (0.01 ~ 0.02 mg/kg) and Korea (0.015 ~ 0.021 mg/kg).38–43 For vegetables, the statistic of Cd concentration of the south and the national are higher than the one reported for Japan (0.02 mg/kg), Korea (0.02 mg/kg), Norway (0.0069 ~ 0.012 mg/kg) and the United Kingdom (0.002 ~ 0.012 mg/kg),38–43 but lower than the one reported in Greece (0.05 mg/kg) and Italy (0.06 mg/kg).38 Owing to the different dietary patterns and Cd concentrations in food groups, the dietary Cd intake rate of adults varies between 0.160 and 0.557 μg/(kg BW·day) among different regions in China. They are all below 0.833 μg/(kg BW·day), which is the PTMI rate issued by the Food and Agriculture Organization (FAO) and the WHO in 2010 (ref. 44). However, compared with 0.365 μg/(kg BW·day), the tolerable value issued to protect children, vegetarians and people living in heavily contaminated regions by the European Food Safety Authority (EFSA) in 2009 and 2011 (refs 45, 46), the average Cd intake rate at the national level, as well as the value of the people in the south and Shanghai are higher. Meanwhile, the average value for the whole population in the present study is somewhat higher than the values reported by Gao et al.26 based on the total dietary surveys carried on in 1990 (13.8 μg/day), 1992 (19.4 μg/day) and 2000 (22.2 μg/day), indicating that the deterioration of food quality due to the contamination occurred during the industrial development process. Compared with the intake rates of people in Korea (14.41 μg/day), the United States (11.5 ~ 14.2 μg/day), Belgium (16.3 μg/day), Denmark (16 μg/day), Finland (16 μg/ day), France (10.6 μg/day), Norway (15.8 μg/day), Germany (19.2 μg/day), the Netherlands (19.3 μg/day), the UK (12.1 μg/day), Portugal (16.5 μg/day) and Canada (16.06 μg/day), the average intake rate of the adults in China is higher. However, in contrast with Japan (25 ~ 30 μg/day), Italy (20.2 μg/day), Ireland (25.1 μg/day) and New Zealand (16.3 μg/day), the difference is negligible.38,43,47–51 The breakdown of Cd contributions from different food groups for adults shows that vegetables are the most important contributor, while the contributions from pulses, eggs, milk and fruits (except Beijing) are negligible. In comparison, rice, vegetables, shellfish, mollusk, seaweed and meat account for 88.6% of the Cd intake for people in Korea.43 Among them, the contribution from rice is the greatest (25.2%) (rice intake rate is also the highest), followed by shellfish (14.5%) and mollusk (11.7%), which



438 also have high Cd concentrations, up to 0.5 and 0.2 mg/kg, respectively.43 For the whole population in Ireland, the dominant contributor is vegetables, up to 70%, followed by flour (23.6%).48

Figure 2.

However, in New Zealand it was reported that for the male adults older than 25 years, the main contributor was oyster (26%), potato (26%) and bread (10%), contributing 62% to the total intake.48

Cd intake from different food types.




439 Table 3.

Soil screening values of Cd (mg/kg)

Land use

Dietary exposure not considered

Dietary exposure considered

National North South Beijing Shanghai Commercial Residential

825 83

458 −1

627 /

365 /

693 /

489 /

In Italy, the key contributors for adults were flour (29%), vegetables (26%) and aquatic products (24%),52 while for adults in France, flour contributed 26%, followed by vegetables (19%), fish and aquatic products (13%) and potato (8%).52 The calculated dietary Cd intake rate of children at the national level was 1.007 μg/(kg BW·day) in this study, which is much higher than the allowable doses issued by the WHO and EFSA. The critical contributor was vegetables, followed by rice, flour, meat and aquatic products, which was similar to that of the adults at the national level. Compared with the daily intake rates reported for children in Germany (0.737 μg/kg BW·day), Australia (0.833 μg/kg BW·day), Denmark (0.31 μg/kg BW·day), New Zealand (0.383 μg/kg BW·day) and the Netherlands (0.32 μg/kg BW·day),51,53–56 the value for children in China is much higher. The main contributors were different among these countries, which may due to the different dietary patterns and Cd concentrations in each food group. In Germany, the main contributors were fruits and vegetables, accounting for 37.1% of the total intake, followed by cereals and bakery goods and beverages.53 In Australia, vegetables accounted for > 50%, followed by cereals and other grainbased foods,54 which was similar to that of China. In contrast with Australia, cereal and cereal products were the predominant contributor to the dietary exposure of children in Denmark,55 accounting for nearly 50%, followed by vegetables. In New Zealand, potatoes were the main source of dietary Cd exposure, accounting for nearly 30%,51 followed by bread. In the Netherlands, wheat was the most important contributor to Cd exposure, accounting for nearly 50%, followed by wheat.56 The present results show that when the dietary exposure to Cd is considered, the SSVs will be reduced for the commercial scenario in which adults were considered as the sensitive risk receptors, especially for the south, where the intake rate was the highest. Therefore, it can be concluded that the hazard will be underestimated and some potential contaminated sites would be screened out as clean sites using screening values without considering dietary exposure. However, compared with the screening value (150 mg/kg)57 issued by the BEPB, assuming the acceptable hazard index to be 0.2, the derived values in this study are much higher. Furthermore, due to the different dietary intake rates among different regions, the difference among the derived values is very large. Under the same exposure scenario, the screening value for the south is the lowest, followed by the value for the whole population, Shanghai and the north, and the value for Beijing is the highest. The derived value for the north is 1.4 times higher than the national average value, while the one for the south is 1.3 times lower. The difference between the values of Beijing and the north (where Beijing is located) is negligible, while the value of Shanghai is 1.3 times higher than the value of the south where it is located. The above differences indicate that if the average Cd intake rate for the whole population is to be used to derive the national SSV, the hazard for the people in the north, Beijing and Shanghai will be overestimated, but for people in other southern cities other than Shanghai, the hazard will be underestimated. © 2015 Nature America, Inc.

Compared with the industrial screening value (1300 mg/kg) issued by New Zealand,11 the results in this study are much lower. The difference is because the dietary intake rate for people in New Zealand (18.2 μg/day) is lower than the values in this study, and because the exposure parameters used by New Zealand are different from the ones used in this study.11 However, in comparison with the soil guideline value of the United Kingdom (230 mg/kg) for the commercial scenario, the values in the study are much higher because the tolerable daily intake (TDI) rate of Cd in the United Kingdom was set to 0.36 μg/(kg BW·day) and the allowed exposure dose from soil was set to half that of the TDI.58 Compared with the regional screening level (100 mg/kg) issued by the US EPA,59 the results in this study are much higher, because its acceptable hazard index is assumed to be 0.1, assuming that 10 toxic elements will do harm to the same organ simultaneously. For the residential scenario in which children were the sensitive risk receptors, the screening value derived theoretically was a negative value, meaning that the dietary Cd exposure of children exceeded the maximum tolerable dose, and that the Cd screening value of Cd should be set to 0 or its concentration in soil should be remediated to 0 if no other risk control measures are taken. However, this may not be rational as the background concentration of Cd in China is 0.02–0.33 mg/kg20 and lowering the Cd concentration in soil to 0 is not cost effective and is not practical from a technical perspective. In this situation, except for reducing the dietary Cd exposure of children through improving food quality or changing dietary patterns, the method used to derive the Cd soil guideline value in the United Kingdom may be a good benchmark on which to set the allowable dose through soil exposure to be 50% of the total maximum allowable intake rates if the non-soil exposure dose was more than a half,58 and the screening value for residential scenario will be 41 mg/kg if this method is applied. However, in China the exact value assigned to the allowed exposure dose through soil should be assessed comprehensively based on our current food quality so that the exact Cd exposure status of children and the economic implications can be assessed thoroughly. CONCLUSIONS The average Cd intake rates for adults in different regions of China are all below the PTMI issued by the FAO/WHO, and range from 0.160 μg/(kg BW·day) and 0.557 μg/(kg BW·day). However, the mean values of the nation, the south and Shanghai exceed the values issued by the EFSA to protect children, vegetarians and people living in heavily contaminated regions. The average intake rate of the whole nation is higher than the values of the north, Beijing and Shanghai, but it is lower than the value of the south. Vegetables are the critical contributor to Cd intake, followed by rice, flour, meats and aquatic products. The screening values for the commercial land-use scenario will be reduced by 16–55.6%, meaning that the hazard will be underestimated if dietary exposure is not considered, and deriving a screening value for each specific area based on its corresponding exposure characteristics of Cd is more appropriate due to the large variation among different regions. The Cd concentration in vegetables is the most sensitive factor in calculating the intake rate, followed by Cd concentrations in rice and aquatic products, and the intake rate of flour, indicating that more attention should be given to these parameters in future total diet surveys. The screening value for the residential land-use scenario derived theoretically based on the whole dietary exposure status of children was a negative value, meaning that the dietary intake rate exceeded the maximum tolerable dose. The method used in the United Kingdom can be applied to derive screening values for the residential scenario based on comprehensive assessment of the exact dietary Cd exposure status of children in China as well as the economic impact of each decision.



440

Figure 3. Sensitivity analysis contribution to variance (%). AP, aquatic product; C, Cd concentration; EG, egg; FL, flour; FR, fruit; IR, intake rate; MI, milk; ME, meat; PL, pulses; RI, rice; VE, vegetable.




441 CONFLICT OF INTEREST The authors declare no conflict of interest.

ACKNOWLEDGEMENTS This work was financially supported by the Beijing Municipal Science and Technology Commission (D08040000360000) and an international cooperative project between the Beijing Municipal Environmental Protection Bureau and the Italian Ministry for Environment, Land and Sea.

REFERENCES 1 MEPPRC. (Ministry of Environmental Protection of the People’s Republic of China). Provisional Rules for the Environmental Management of Contaminated Sites (Draft) Beijing: MEPPRC 2009 (in Chinese). 2 MEPPRC. (Ministry of Environmental Protection of the People’s Republic of China) Technical Guideline for Risk Assessment of Contaminated Sites 2014, HJ25.3-2014 Beijing: MEPPRC (in Chinese). 3 JSEPD (Environmental Protection Department of Jiangsu Province) Notes on Pollution Prevention and Control on Industrial Contaminated Sites 2014 Nanjing: JSEPD (in Chinese). 4 SYEPB (Shenyang Environmental Protection Bureau) Notes on Contaminated Sites Pollution Control and Remediation Shengyang: SYEPB 2007 (in Chinese). 5 CQMPG (Chongqing Municipal People’s Government) Notes on Contaminated Sites Remediation within Chongqing Municipality 2008 Chongqing: CQMPG (in Chinese). 6 PGZJ (The People’s Government of Zhejiang Province) Notes on Supervision on the Redevelopment of Industrial Sites 2013 Hangzhou: PGZJ (in Chinese). 7 Jiang L, Li LX, Huang HL, Wang JL, Wang Y, Zhang LN et al. Environmental site assessment guideline 2009 DB11/T 656-2009 Beijing: Beijing Municipal Quality and Technology Supervision Bureau (in Chinese). 8 Jiang L, Yu LF, Ding Q, Li FS Technical Research on Environmental Risk Management on Contaminated Sites of POPs Pesticides. China Environmental Science Press: Beijing, China, 2012 (in Chinese). 9 Jiang L, Zhong MS, Liang J, Yao JJ, Xia TX, Fan YL. Application and benefit assessment of tiered health risk assessment approach on site contaminated by benzene. Environ Sci 2013; 34: 1034–1043 (in Chinese). 10 ASTM. Standard Guild for Risk-Based Reaction Action 2004 (E2081-00). Washington DC: American Society for Testing and Materials. 11 MfE (Ministry for Environment). Guidelines for Assessing and Managing Petroleum Contaminated Sites in New Zealand 2011. Wellington: Ministry for Environment. 12 Zhu PF, Ning P, Zeng XD, Wang HJ, Zhao R, He B. Soil pollution and heavy metals hyper-accumulators around two sites contaminated by smelting. Safety Environ Engineer 2006; 13: 48–51 (in Chinese). 13 Zhang Y, Liu JJ, Liang CH, Chen XZ. Study on pollution status of heavy metal in soil from factory district of a former smelter in Shenyang City. J Anhui Agric Sci 2008; 36: 6481–6483 (in Chinese). 14 Zeng YM, Ye ZR. Characteristics and assessment of soil contamination by heavy metals around zinc smelting plant of Huixian. Sci Technol Inform 2010; 7: 346–347 (in Chinese). 15 Du P. Distribution of Heavy Metals and Their Chemical Speciation in Soil Around a Pb-Zn Smelter. Chinese Research Academy of Environmental Science: Beijing, China, 2007 (in Chinese). 16 Jiang Y, Zhan XW, Sun LD, Xiao RM, Huang ZL. The formation distribution of Zn, Cu and Pb and Cd in the soil nearby the smelter. Guangdong Chem Ind 2011; 38: 110–112 (in Chinese). 17 Li XH, Tang ZL. Concentrations of cadmium, copper, lead, zinc in soil around smelting plant of Baiyin Region, Gansu Province. Geotech Invest Survey 2007; 5: 23–27 (in Chinese). 18 Niu YL Study on Heavy Metal Dynamic Distribution Characteristic and its Pollution Warnimg. Hunan University: Changsha, China, 2007 (in Chinese). 19 Xu ZJ, Wu CH, Liu F, Qiu XY. Study on combined pollution character of heavy metals in soils around a typical Lead-Zinc smeltery. J Hum Univ Sci Technol 2007; 22: 111–114 (in Chinese). 20 Zhao XJ, Lu SJ, Xu RJ, Li BL, Wu GP, Wui FS. Soil heavy metal Cadmium standard limit and range of background value research. Environ Sci 2014; 35: 1491–1496 (in Chinese). 21 Yu H, Wang JL, Fang W, Yuan JG, Yang ZY. Cadmium accumulation in different rice cultivars and screening for pollution-safe cultivars of rice. Sci Total Environ 2006; 370: 302–309. 22 Yuan XP, Wang J, Shang YE, Sun BG. Health risk assessment of cadmium via dietary intake by adults in China. J Sci Food Agric 2014; 94: 373–380.


23 Liu H, Wu CF, Lu Y, Fang YM, Gao WW, Xin ZH. Assessment on the dietary exposure of lead and cadmium in Shanghai residents. Chinese J Food Hygiene 2011; 23: 218–223 (in Chinese). 24 Satarug S, Garrett SH, Sens MA, Sens DA. Cadmium, environmental exposure, and health outcomes. Environ Health Perspect 2010; 118: 182–190. 25 Huang LC, Tang J, Zhang RH, Gu F, Fang YQ, Zhang HX et al. Exposure assessment on dietary lead and cadmium of residents in three areas of Zhejiang Province from 2009 to 2010. Chinese J Prevent Med 2012; 46: 42–46 (in Chinese). 26 Gao JQ, Li XW, Zhao JL. 2000 Chinese total diet study—the dietary lead and cadmium intakes. J Hygiene Res 2006; 35: 750–754 (in Chinese). 27 CRAES (Chinese Research Academy of Environmental Science). Technical Document on Deriving Soil Screening Values of Beijing. Beijing Environmental Protection Bureau: Beijing, 2011. 28 Jin SG. The Tenth Nutrition and Health Survey Report about Chinese Residents: Dataset about Nutrition and Health in 2002. People’s Medical Publishing House: Beijing, 2008 (in Chinese). 29 Han YF, Zhang JG, Zhang ZS. Regional Geography of China. Guangdong Higher Education Press: Guangzhou, China, 2009 (in Chinese). 30 Li G, Sun GX, Williams PN, Nunes L, Zhu YG. Inorganic arsenic in Chinese food and its cancer risk. Environ Internat 2011; 37: 1219–1225. 31 MoH (Ministry of Health of the People’s Republic of China). Contaminants Limit in f#ood 2012 GB 2762-2012 (in Chinese). Beijing: MoH 32 Williams PN, Price AH, Raab A, Hossain SA, Feldmann J, Meharg AA. Variation in arsenic speciation and concentration in paddy rice related to dietary exposure. Environ Sci Technol 2005; 39: 5531–5540. 33 Khan SI, Ahmed AK, Mottashir Y, Mohammad R, Mahfuzar H, Samar K et al. Arsenic and Cadmium in Food-chain in Bangladesh--An Exploratory Study. J Health Population Nutr 2010; 28: 578–584. 34 Gemma P, Roser MC, Juan ML, Josel D. Effects of Various cooking processes on the concentrations of arsenic, cadmium, mercury, and lead in foods. J Agric Food Chem 2008; 56: 11262–11269. 35 Zhang HL, Li XP, Qi JY, Chen YT, Li BJ. Primary research on health risk assessment of heavy metals in the surrounding soil and air of a municipal solid waste incinerator (MSWI), south China. J Agro Environ Sci 2013; 32: 1670–1676 (in Chinese). 36 MfE (Ministry for Environment). Methodology for Deriving Soil Guideline Values Protective of Human Health 2010 Wellington: Ministry for Environment. 37 John AC, Richard LB, Thomas EM Software Guidance Manual-RBCA Tool Kit for Chemical Releases. GSI Environmental Inc: Houston, 2012. 38 EC (European Commission). Assessment of the Dietary Exposure to Arsenic, Cadmium, Lead and Mercury of the Population of the EU Member States 2004. Brussels: European Commission. 39 Yuriko K, Testuo N, Nami K, Takamoto U, Kazaguki O. Cadmium concen tration in current Japanese foods and beverages. J Occup Health 2002; 44: 240–247. 40 Health Canada. Trace Element Concentrations in 2007 Total Dietary Study Composites. Available at http://www.hc-sc.gc.ca/fn-an/surveill/total-diet/concentration/ conc-food-alim_2007-eng.php. 2007 (accessed 7 March 2014). 41 Tsukahara T, Ezaki T, Moriguchi J, Furuki K, Shimbo S, Matsuda-Inoguchi N et al. Rice as the most influential source of cadmium intake among general Japanese population. Sci Total Environ 2003; 305: 41–51. 42 Zazouli MA, Shokrazdeh M, Izanloo SF. Cadmium content in rice and its daily intake in Ghaemshahr region of Iran. Afr J Biotechnol 2008; 7: 3686–3689. 43 Kim M, Wolt JD. Cadmium exposure in the South Korean population: implications of input assumptions for deterministic dietary assessment. Hum Ecol Risk Assess 2008; 14: 835–850. 44 FAO/WHO (Food and Agriculture Organization/World Health Organization). Working Document for Information and Use in Discussions Related to Contaminants and Toxins in the GSCTFF (General Standard for Contaminants and Toxins in Food and Feed) 2011. Geneva: FAO/WHO. 45 EFSA (European Food Safety Authority). Scientific Opinion of the Panel on Contaminants in the Food Chain on a Request from the European Commission on Cadmium in Food 2009. 46 EFSA (European Food Safety Authority). Statement on Tolerable Weekly Intake for Cadmium 2011. 47 Egan SK, Bolger PM, Carringto CD. Update of US FDA’s Total Diet Study food list and diets. J Expos Sci Environ Epidemiol 2007; 17: 573–582. 48 FSAI (Food safety authority of Ireland). Report on a Total Diet Study Carried Out by the Food Safety Authority of Ireland in the Period 2001–2005, Dublin:Food safety authority of Ireland 2011. 49 Health Canada. Dietary Intake of Contaminants and Other Chemicals for Different Age-Sex Groups of Canadians. Available at http://www.hc-sc.gc.ca/fn-an/surveil/ total-diet/intake-apport/index-eng.php (accessed 7 March 2014).



442 50 Ikeda M, Ezaki T, Tsukahara T, Moriguchi J. Dietary cadmium intake in polluted and non-polluted areas in Japan in the past and in the present. Int Arch Occup Environ Health 2004; 77: 227–234. 51 MAF (Ministry of Agriculture and Forestry). 2009 New Zealand Total Diet StudyAgricultural Compound Residual, Selected Contaminant and Nutrient Elements 2011. Wellington: Ministry of agriculture and Forestry. 52 EFSA (European Food Safety Authority). Cadmium Dietary Exposure in the European Population 2012. 53 DGHCP (Directorate-General Health and Consumer Protection). Assessment of the Dietary Exposure to Arsenic, Cadmium, Lead and Mercury of the Population of the EU Member States 2004. 54 FSANZ (Food Standards Australia New Zealand). The 23rd Australian Total Diet Study 2011.

55 NFI (National Food Institute). Chemical Contaminants 2004–2011, 2013. 56 RIVM (The Netherlands National Institute of Public Health and the Environment). Dietary Intake of Heavy Metals (Cadmium, Lead and Mercury) by the Dutch Population 2003. 57 Zhou YY, Li FS, Li LX, Huang HL, Cao YZ, Yan ZG et al. Screening Levels for Soil Environmental Risk Assessment of Sites. Beijing Municipal Quality and Technology Supervision Bureau: Beijing, 2011 DB11/T 811-2011. (in Chinese). 58 EA (Environmental Agency). Soil Guideline Values for Cadmium in Soil 2009. London: EA. 59 US EPA. Regional Screening Level (RSL) Industrial Soil Table (TR = 1E-6, HQ = 0.1). Available at http://www.epa.gov/reg3hwmd/risk/human/rb-concentration_table/ Generic_Tables/docs/indsoil_sl_table_01run_NOV2013.pdf. accessed 7 March 2014.

Supplementary Information accompanies the paper on the Journal of Exposure Science and Environmental Epidemiology website (http:// www.nature.com/jes)

